Changes in trophic niches of oribatid mites with transformation of tropical rainforest systems - from rainforest into rubber and oil palm plantations in Sumatra, Indonesia

(1)

C HANGES IN TROPHIC NICHES OF ORIBATID MITES WITH TRANSFORMATION OF TROPICAL RAINFOREST SYSTEMS – FROM

RAINFOREST INTO RUBBER AND OIL PALM PLANTATIONS IN SUMATRA , I NDONESIA

Dissertation

For the award of the degree

“Doctor rerum naturalium” (Dr.rer.nat.) of the Georg-August-Universität Göttingen

within the doctoral program Biodiversity and Ecology of the Georg-August University School of Science (GAUSS)

submitted by

M.Sc. Alena Krause

from Gifhorn Göttingen, 2020

(2)

Thesis Committee

Stefan Scheu / Animal Ecology, J.F. Blumenbach Institute of Zoology and Anthropology, Göttingen (Name of Department / Research Group, Institution)

Marko Rohlfs / Institute of Ecology, Population and Evolutionary Biology, Bremen (Name of Department / Research Group, Institution)

Mark Maraun / Animal Ecology, J.F. Blumenbach Institute of Zoology and Anthropology, Göttingen (Name of Department / Research Group, Institution)

Members of the Examination Board

Reviewer Stefan Scheu / Animal Ecology, J.F. Blumenbach Institute of Zoology and Anthropology, Göttingen

(Name of Department / Research Group, Institution)

Second Reviewer Marko Rohlfs / Institute of Ecology, Population and Evolutionary Biology, Bremen (Name of Department / Research Group, Institution)

Additional Reviewer (if applicable)

(Name of Department / Research Group, Institution) Further member of the Examination Board

Mark Maraun / Animal Ecology, J.F. Blumenbach Institute of Zoology and Anthropology, Göttingen (Name of Department / Research Group, Institution)

Christoph Bleidorn / Animal Evolution and Biodiversity / J.F. Blumenbach Institute of Zoology and Anthropology, Göttingen

Holger Kreft / Biodiversity, macroecology, biogeography / Faculty of Forest Sciences and Forest Ecology, Göttingen

Klaus Hövemeyer / Animal Ecology, J.F. Blumenbach Institute of Zoology and Anthropology, Göttingen (Name of Department / Research Group, Institution)

Date of the oral examination: 11.05.2020

(3)

“Soil provides the foundation of human existence, […].” from “Soil Fauna Assemblages, Global to Local Scales” from Uffe N. Nielsen

(4)

(5)

Content

Summary ... 1

General Introduction ... Land-use change and different land-use systems ... 5

Soil communities ... 7

Macrofauna ... 9

Oribatid mites ... 9

Trophic ecology ... 10

Trophic niches ... 10

Trophic plasticity ... 13

Stable isotopes ... 14

Management of oil palm plantations and the ‘Biodiversity Enrichment Experiment‘ ... 17

Study site ... 20

Objectives and chapter outline ... 24

References ... 27

Shift in trophic niches of soil microarthropods with conversion of tropical rainforest into plantations as indicted by stable isotopes (¹⁵N, ¹³C) ... Abstract ... 40

Introduction ... 41

Material and Methods ... 44

Study sites ... 44

Sampling, extraction and species determination ... 45

Stable isotope analysis ... 46

Statistical analysis ... 47

Results ... 48

Discussion ... 51

Trophic niches of species ... 51

Shift in trophic niches with land use ... 53

Conclusions ... 56

Acknowledgments ... 56

Funding ... 56

Field work permissions ... 57

Sample collection and determination ... 57

References ... 57

Appendix ... 64

(6)

Content

Variation in community level trophic niches of soil microarthropods with conversion of tropical rainforest into plantations systems as indicted by stable isotopes (¹⁵N, ¹³C) ...

Abstract ... 76

Introduction ... 78

Stable isotope analysis ... 83

Results ... 88

Trophic structure ... 88

One-dimensional metrics ... 89

Multidimensional metrics ... 91

Discussion ... 93

One-dimensional metrics ... 93

Multidimensional metrics ... 97

Conclusions ... 98

Funding ... 99

References ... 100

Appendix ... 106

Response of soil animal communities to tree diversity enrichment of oil palm plantations ... 1

Introduction ... 132

Results ... 138

Abundance of total macro- and mesofauna ... 138

Richness of macro- and mesofauna taxa... 140

Community structure ... 142

Discussion ... 144

Abundance of macro- and mesofauna ... 144

Richness of macro- and mesofauna taxa... 145

(7)

Content

Community structure ... 147

Conclusions ... 148

Funding ... 149

References ... 150

Appendix ... 158

General Discussion ... 1

Conclusions ... 171

References ... 172

Acknowledgments ... 177

List of publications ... 180

Thesis declaration ... 181

Declaration of the author’s own contribution to manuscripts with multiple authors... 181

Plagiarism declaration ... 183 Poster and oral presentation ... Fehler! Textmarke nicht definiert.

(8)

(9)

Summary

1

Summary

During the last decades especially tropical regions suffered from degradation as well as transformation of landscapes into different land-use systems. Logged rainforest sites in Southeast Asia are often transformed into cash crop monocultures, especially oil palm and acacia plantations. Such transformation processes may threaten the functioning of ecosystems with the worldwide highest biodiversity and endemism. Effects of this transformation and degradation have mainly been studied for aboveground organisms whereas effects on the functioning and composition of soil invertebrate communities are little studied.

This thesis focuses on the effects of land-use transformation along a land-use gradient, i.e.

from secondary rainforest to plantations (jungle rubber, rubber and oil palm monoculture), on microarthropod communities, using oribatid mites as model organisms.

Further, I investigated the effect of management strategies within oil palm plantations on macro- and mesofauna soil animals. The field studies were conducted within the interdisciplinary project “Ecological and socioeconomic functions of tropical lowland rainforest transformation systems (Sumatra, Indonesia)” (EFForTS), established in Jambi Province, southwest Sumatra (Indonesia) in 2013.

In the first study, presented in Chapter 2, we investigated shifts in trophic niches of six soil-living oribatid mite species and their possible trophic plasticity with the conversion of lowland secondary rainforest into plantation systems (jungle rubber, rubber and oil palm monoculture plantations) in two regions of southwest Sumatra, Indonesia. Therefore, stable isotope ratios (¹³C/¹²C and ¹⁵N/¹⁴N) of single oribatid mite individuals were measured and, subsequently, we calculated shifts in stable isotope niches with changes in

(10)

Summary

2

land-use systems. On the basis of significant changes in stable isotope ratios in three of six studied oribatid mite species this study demonstrated that these species are able to shift their trophic niche in land-use transformation systems. Those shifts were either due to changes in trophic level (indicated by δ¹⁵N values) or due to changes in the use of basal resources (indicated by δ¹³C values) or both. Notably, the shifts were most pronounced between more natural systems (secondary rainforest and jungle rubber) on one side and monoculture plantation systems (rubber and oil palm plantations) on the other side;

thereby indicating that the shifts were related to land-use intensity.

In the second study, presented in Chapter 3, we investigated shifts in community-level trophic niches of soil-living oribatid mites with the conversion of lowland secondary rainforest into plantation systems (jungle rubber, rubber and oil palm monoculture plantations) in two regions of southwest Sumatra, Indonesia. Therefore, stable isotope ratios (¹³C/¹²C and ¹⁵N/¹⁴N) of pooled oribatid mite species were measured, and subsequently, we calculated shifts in community-level trophic niche with transformation of land-use systems. This study demonstrated that the community-level trophic niche of oribatid mites is larger in rainforests than in plantation systems, suggesting that the conversion of rainforest into plantation systems is associated with reduced availability of litter resources. Results of this study further demonstrated that community-level trophic niches in rainforest and jungle rubber are separated from those in monoculture plantation systems, indicating again that the trophic niche of oribatid mite communities shifts markedly with land-use change. Additionally, ¹⁵N/¹⁴N ratios of oribatid mite communities indicated that the diet of microarthropods shifts towards predation and/or scavenging with changing land-use systems. This may be due to the limited amount of litter and its

(11)

Summary

3

low quality in rubber and oil palm plantations. Further, exceptionally low ¹³C/¹²C ratios of oribatid mite communities in rubber plantations suggest that certain oribatid mite species in these land-use systems use resources which are lacking in the other studied ecosystems.

Oribatid mite communities in plantation systems present an unusual high functional richness and uniqueness compared to natural systems.

The results of the first two studies demonstrated that soil-living oribatid mite species are able to adapt to changing land-use systems and do not suffer to the same extent from these changes as many aboveground species. The third experiment, presented in Chapter 4, focused on investigating the effect of ‘tree islands’ of different size (5 x 5, 10 x 10, 20 x 20 and 40 x 40 m) and diversity level of planted native trees (0, 1, 2, 3 and 6) within oil palm plantations. Here we investigated the response of meso- and macrofauna species to the establishment of ‘tree islands’ three years after the experiment started. Neither the different diversity levels of native tree species nor the plot size significantly affected the abundance of soil invertebrate taxa. However, richness of soil invertebrate taxa was positively affected in ‘tree islands’ of diversity level 2. The result demonstrated that the diversity and abundance of plant communities little affect the structure and diversity of soil invertebrates three years after establishment suggesting soil invertebrates respond with a pronounced time lag to the experimental manipulations. Overall, by investigating the trophic ecology of oribatid mites and their response to changes in land-use systems the results of this thesis improved the understanding of how soil communities and individual species respond to the conversion of rainforest into intensively managed agricultural systems.

(12)

(13)

Chapter I

General Introduction

(14)

(15)

General Introduction

5

Land-use change and different land-use systems

The worldwide rapidly growing human population is leading to a rising need for food, fuel and fiber, and therefore transformation as well as degradation of landscapes is increasing (Dirzo and Raven, 2003; Foley et al., 2005; Gibbs et al., 2010; Newbold et al., 2015).It is estimated, that the worldwide population size will increase to 9.7 billion by 2050 (UN, 2015) leading to an increase in the demand for food by 70 % (Godfray et al., 2010). These demands lead to high pressure on ecosystems worldwide, leading to a higher conversion of natural ecosystems into plantations, with more pressure on the production and yield of those agricultural systems (Godfray et al., 2010; Lambin and Meyfroidt, 2011; Tscharntke et al., 2012). Human activity strongly impacts natural ecosystems, directly e.g., via building infrastructure and houses, as well as indirectly e.g., via climate change and nutrient deposition (DeFries et al., 2004; Foley et al., 2005, 2011). About 40 % of the terrestrial surface has been transformed into agricultural systems, with an increasing proportion being degraded resulting in habitat loss and desertification e.g., due to erosion, due to construction of infrastructure and additionally due to human behavior (Bridges and Oldeman, 1999; Reynolds et al., 2007; Foley et al., 2011; Pavao-Zuckerman and Sookhdeo, 2017; Francini et al., 2018). This is mainly caused by high levels of fertilizer application and atmospheric deposition. Additionally, external input of nitrogen and phosphorus has been increasing since the 19^th centuries (Peñuelas et al., 2012).

In the last decades, especially tropical regions suffered from degradation and transformation into land-use systems, such as oil palm or rubber plantations (Sodhi et al., 2010; Wilcove et al., 2013; Meijide et al., 2018). In South East Asia these transformations are threatening ecosystems with the highest biodiversity and endemism worldwide

(16)

6

(Jones, 2013). Since the mid-20^th century rainforests in Southeast Asia have been logged, often followed by the transformation of logged sites into cash crop monocultures, such as rubber, oil palm and acacia plantations (Koh and Wilcove, 2008; Wilcove and Koh, 2010;

Drescher et al., 2016). The process of expanding as well as intensification of agricultural landscapes poses the greatest threat to biodiversity (Tilman et al., 2001; Donald, 2004;

Green et al., 2005). In my thesis, a gradient of different land-use systems was studied.

Rainforest sites were represented by ‘primary degraded forest’, classified after Margono (2014). Jungle rubber land-use systems were rubber agroforests systems (Hevea brasiliensis) that resemble secondary rainforest, where naturally occurring species of different trees were included (Beukema et al., 2007). Rubber monoculture plantations exclusively include rubber trees (Hevea brasiliensis), and oil palm monoculture plantations exclusively include oil palm trees (Elaeis guineensis) (Drescher et al., 2016).

One of the agricultural land-use systems that is rapidly increasing are vegetable oils (Clay, 2013), with oil palm as one of the most quickly expanding crops (Carter et al., 2007;

Fitzherbert et al., 2008). Additionally, biofuel markets and rising food demand in the European Union as well as in Indonesia, India and China result in increasing global oil palm production, currently by about 9 % each year (European Comission, 2006; Clay, 2013), with Malaysia and Indonesia as the main producers of palm oil (Koh and Wilcove, 2007). Palm oil belongs to one of the versatile oils, which has not only many different functions and therefore is widely used but it is also one of the most efficient crops worldwide (Zimmer, 2010; Ashraf et al., 2018). Moreover, it is the crop which produces the highest yield per land area (Zimmer, 2010; Ashraf et al., 2018).

(17)

7

Transformation of rainforest into agricultural systems strongly increased in Indonesia.

Commercial oil palm cultivation in Indonesia started in 1911, with Sumatra as starting point (Abdullah and Nobukazu, 2007; Corley et al., 2008). After the 1980s oil palm plantations were also established in other parts of Indonesia (Abdullah and Nobukazu, 2007; Corley et al., 2008). Oil palm as well as rubber plantations often were established on rainforest sites which were already logged or degraded by fire (Curran et al., 2004;

Dennis et al., 2005; Fitzherbert et al., 2008; Drescher et al., 2016). Nevertheless, conversion of rainforest into oil palm plantations may account for 16 % of recent deforestation in Indonesia (Fitzherbert et al., 2008), whereas the conversion of rainforest into rubber plantations and therefore the production of natural rubber has increased more than 50 % since 2000 (Ahrends et al., 2015).

In 2012, 0.84 million hectares rainforest were converted into agricultural systems in Indonesia, the highest rate worldwide (Margono et al., 2014; Drescher et al., 2016). One of the highest losses of primary forest occurred in Sumatra (Indonesia), with 0.40 million hectares per year between 2009 and 2011 (Laumonier et al., 2010; Miettinen et al., 2011;

Margono et al., 2014). Oil palm plantations are known to hold less than half as many vertebrate species as primary rainforest (Danielsen et al., 2009). However, the effects of conversion of rainforest into plantation systems have been rarely studied for belowground arthropods (Newbold et al., 2015).

Soil communities

Agricultural production essentially depends on soil, and soils therefore are important for human welfare, e.g. food, fiber and fuel production (Nielsen, 2019). In fact, one of the key

(18)

8

factors for the survival of humankind relies on soils and soil processes which are based on the activity of soil biota. However, until today understanding of the structure of soil communities and the functioning of soil systems still is limited.

Up to 90 % of the primary production of plants enters the soil system as leaf/wood detritus and rhizodeposits (McNaughton et al., 1989; Bardgett, 2005). Therefore, decomposition, together with primary production, is the most important process for terrestrial ecosystems. The presence or absence of specific soil animal species can modify the structure and functioning of soil systems, such as the turnover of organic matter and nutrient cycling (Bardgett, 2005; Nielsen et al., 2015). Additionally, certain functional types of soil fauna may enhance soil functioning or even modify soil food webs (Brussaard et al., 2007). Soil arthropods are part of any soil but the abundance and diversity varies strongly between different ecosystems, even within small spatial scales (Ettema and Wardle, 2002). This high spatial heterogeneity in the structure of soil communities is likely due to variations in biotic as well as abiotic factors, e.g. climate and litter type (Coûteaux, Marie-Madeleine Bottner and Berg, 1995; Wardle et al., 2006; Berg and McClaugherty, 2008). Soil fauna is highly diverse and can be divided into functional size classes of micro- , meso- and macrofauna (Swift et al., 1979). The trophic differentiation of micro- and macrofauna species has been long accepted and recent evidence based on stable isotope analysis underlined these assumptions (Potapov et al., 2019), whereas mesofauna taxa usually were taken as uniform trophic guild (De Ruiter et al., 1993; Berg and Bengtsson, 2007; Moore and de Ruiter, 2012). Contrasting this assumption, recent studies based on stable isotopes documented a variety of trophic niches and trophic levels within major groups of mesofauna (Schneider et al., 2004; Chahartaghi et al., 2005; Maraun et al., 2011;

(19)

9

Klarner et al., 2013). As Potapov et al. (2019) stated, different species within the same taxonomic group can belong to different trophic levels and may therefore provide different ecosystem services. The most abundant taxa for soil mesofauna are Collembola and Acari (Petersen and Luxton, 1982). Both groups represent a wide range of different life history traits, trophic positions, and therefore are likely to affect ecosystem functions in a variety of ways (Scheu, 2002; Schneider et al., 2004; Nielsen, 2019).

Macrofauna

Macro- and megafauna, due to their larger size, have different and more pronounced effects on ecosystems (Lal, 1988; Folgarait, 1998; Migge-Kleian et al., 2006). They play an important role for litter fragmentation as well as displacement, produce large amounts of faecal pellets and can enhance decomposition processes (David, 2014). Many macro- and megafauna species are considered ecosystem engineers, since they can modify the environment strongly, e.g. by feeding, burrowing, and the production of faecal pellets (Bonachela et al., 2015; Parr et al., 2016; Ashton et al., 2019). Additionally, macro- and megafauna produce lasting imprints on the environment, including the change of soil structure as well as organic matter distribution, and thus influence soil properties, soil biological assemblages, element cycling and ecosystem functioning more than other soil fauna (Migge-Kleian et al., 2006; Parr et al., 2016; Ashton et al., 2019; Nielsen, 2019).

Oribatid mites

With about 11,000 described species (Subías et al., 2018), and the true number of species likely exceeding 50,000 (Walter and Proctor, 2013), oribatid mites are the most diverse

(20)

10

soil microarthropods. Oribatid mites colonize a wide range of different habitats, e.g.

temperate to tropical regions, deserts, tundras and aquatic habitats (Krantz et al., 2009).

Densities of oribatid mites can reach up to 200,000 ind./m² in forest soils of temperate regions whereas in tropical regions densities typically are in the range of 30,000-40,000 ind./m² (Maraun and Scheu, 2000; Maraun et al., 2007; Scheu et al., 2008). Oribatid mites are trophically highly diverse and span over about three to four trophic levels, including lichen and algae feeders, fungal feeders, primary and secondary decomposer as well as predators/scavengers (Maraun et al., 2004; Schneider et al., 2004; Illig et al., 2005;

Erdmann et al., 2007). Different life history traits, e.g. low fertility, slow development and long life cycles, leave oribatid mites sensitive to soil conditions and thereby changes in environmental conditions (Behan-Pelletier, 1999). Due to high population density and species richness oribatid mites have been proposed as indicator organisms for soil health (Bedano et al., 2011) and land use (Zhao et al., 2013), reflecting impacts of land use intensification, especially in tropical rainforest systems (Migge-Kleian et al., 2007; Gan et al., 2014; Hasegawa et al., 2014).

Trophic ecology Trophic niches

One of the most important concepts in ecology is the niche concept (Hutchinson, 1959;

Chase and Leibold, 2003). Niche differentiation is considered to be the basis for species co-existence (Tokeshi, 2009). Organisms not only interact as an ecological ‘guild’, e.g.

group of taxa that use the same class of resources in a comparable way (Root, 1967), or due to trophic relations, e.g. predator-prey interactions which can shape the process of

(21)

11

evolution but also contributes to the complexity of communities (Tokeshi, 2009), but also due to the use of habitats (Tokeshi, 2009). Therefore, two types of niches are differentiated, the ‘fundamental niche’ (Hutchinson, 1978) and the realized niche, describing the ‘role’ of the species within a community, focusing mainly on its trophic position (Elton, 1927). The fundamental niche is defined as the niche occupied by species in the absence of competition or other biotic interactions, whereas the realized niche is defined as the niche space occupied in presence of competition and biotic factors. Based on the interactions between species and/or populations different mechanisms are responsible for these interactions or effects on other organisms (Abrams, 1987). Six different interaction types are commonly distinguished, i.e. competition, predation, herbivory, parasitism, diseases and mutualism (Krebs, 1994). One of the important factor is the availability of resources for species, forming the basis of competitive interactions (White, 1993). There are two different types of competition as defined by Birch (1957).

First, resource competition occurs when organisms (from the same or different species) indirectly interact by using the same resources which is scarce. Second, interference competition occurs when organisms directly interact for access to resources. Competition therefore can lead to changes in the population size of the competing species (Lotka- Volterra equation; Lotka, 1925; Volterra, 1926) or lead to one species ‘winning’ and the extinction of the other or co-existing of species, based on the availability of resources (Tilman, 1977, 1986).

There are different niche-related concepts. The first is environmental filtering, which implies that communities are assembled according to similarity of niches (Vellend, 2010;

Kraft et al., 2015). The second is based on competition between species leading to niche

(22)

12

differentiation of species within communities (Macarthur et al., 1967; Violle et al., 2011), implying that the structure of communities evolves with the co-existence of species in a stable environment (Korotkevich et al., 2018). The trophic niche as dimension of the ecological niche implies effects of species on other species within communities, thereby being related to the role within the ecosystem of those species (Leibold, 1995; Chase and Leibold, 2003). Species with a broader trophic niche are predicted to more easily invade existing communities than those with a narrower niche, and to survive disturbances more easily (Bommarco et al., 2010; Dammhahn et al., 2017). There are few studies focusing on changes in trophic-niches in disturbed habitats, most of the conducted studies focused on the response of individual species and not the community (Korotkevich et al., 2018).

Studies focusing on trophic-niche shifts at the community level are mostly based on aquatic systems (di Lascio et al., 2013; Hansen et al., 2018). One of the few studies analyzing shifts in trophic niches of soil invertebrates to changes in land use systems, regarding the response of individual species is Krause et al. (2019), with the results indicating trophic plasticity in oribatid mite species. Further, focusing on the community level Klarner et al. (2013) investigated the trophic structure of Mesostigmata in beech stands in Central Germany and showed that Mesostigmata predominantly feed on secondary decomposer. Notably, closely related taxa often had very different stable isotope values suggesting that trophic niche partitioning allowed the coexisting of morphological similar species (Klarner et al., 2013).

(23)

13

Trophic plasticity

Trophic plasticity allows animals to react to changing environmental conditions, such as global warming, intensified land use or flooding. Organisms with generalist feeding habits are more flexible regarding their diet and therefore may be less affected by changing environmental conditions as compared to organisms with a more narrow diet. Until today, trophic plasticity mostly has been investigated in aquatic taxa, mainly in fish (Bowen and Allanson, 1982; Almeida et al., 2012; Drymon et al., 2012) and snails (Riera, 2010).

Predominantly, these studies focused on changes in morphology and behavior due to changing environmental factors rather than on trophic plasticity. Juvenile Tilapiu mossumbica (Cichlidae. Teleostei) move daily from deep offshore waters to shallow littoral areas for feeding and back (Bowen and Allanson, 1982). This movement is linked to changes in physical and biological features of the littoral environment and therefore varies in time. Moreover, with changes in the lake water level there were also changes in the diet and behavior. Micropterus salmoides (Percomorphaceae; Teleostei) is known as an invasive species and trophic plasticity likely contributes to the success as invasive species (Almeida et al., 2012). Another important factor for trophic plasticity in one species could be regional variation of diet as shown for Rhizoprionodon terraenovae (Carcharhinidae, Selachii) as well as for Hydrobia ulvae (Hydrobiidae, Gastropoda) (Riera,

2010; Drymon et al., 2012). Leal et al. (2015) showed that tropic plasticity can also be influenced by symbiosis.

Only few studies investigated the trophic plasticity in the field in soil organisms and very few considered changes in trophic niches with changes in land use. One of the few field studies existing showed that centipede predators are able to switch their diet from feeding

(24)

14

on secondary decomposer in rainforest to less ¹³C enriched prey in oil palm plantations (Klarner et al., 2017). Results of another study on predation of centipedes showed that the management of different forest types in Germany does not affect the prey spectrum, but it varies with the depth of the litter layer and soil pH (Günther et al., 2014). Another study focused on oribatid mites from temperate systems (Gan et al., 2014), where oribatid mites were found to suffer from environmental changes, since those animals are assumed to be specialized regarding their diet (according to their ¹⁵N and ¹³C signature) and therefore likely to go extinct with changes in environmental conditions. Generally, laboratory studies with oribatid mites suggest that their food preferences are innate and little affected by learning (Brückner et al., 2018).

Stable isotopes

It is difficult to study the trophic interactions of soil animals (Pollierer et al., 2009). One major problem is the structure of the soil. Soil structure, e.g. pore size distribution, water infiltration, water holding capacity and/or chemical characteristic have a direct impact on the abundance and distribution of soil animals (Ducarme et al., 2004; Nielsen et al., 2008;

Nielsen, 2019). Soil animals are often small, displace complex trophic levels and are not easy to identify (Sunderland et al., 2005; Potapov et al., 2019). Further, the feeding behavior of soil animals is also affected by biotic and abiotic soil conditions (Traugott et al., 2013). Models and simulations can provide insights into dynamics and structure of food webs, however, it is important to validate the trophic position of species/taxa with real-world data (Finlay-Doney and Walter, 2012). Recently, significant methodological advantages have been made studying trophic interactions (Traugott et al., 2013). Some of

(25)

15

the important methods for studying trophic interactions are stable isotope and fatty acid analysis which allow to detect resource allocation (Ruess and Chamberlain, 2010;

Boecklen et al., 2011), whereas DNA-based techniques allow to link feeding interactions to taxonomic positions (Gariepy et al., 2007; Symondson, 2012). Advantages of DNA-based methods include for example that multiple individual samples can be pooled in e.g., NGS- based techniques (Deagle et al., 2009) or gut-content analysis (Zaidi et al., 1999), however, there also are disadvantages. One of the most critical one is the high sensitivity of PCR, which may not only detect food and predator DNA but also DNA of contaminations (Traugott et al., 2013). Fatty acid analysis can be used to detect different diet due to specific fatty acid signatures of bacteria, fungi, algae and plants which animals are not able to synthesize (Ruess and Chamberlain, 2010). Fatty acids of consumers to a large extent originate from their diet as they are assimilated and incorporated without major change ('dietary routing'; Ruess and Chamberlain, 2010; Traugott et al., 2013). Nevertheless, fatty acid analysis also has disadvantages, one is the intermediate specificity as well as the metabolic modification of fatty acid signatures in the consumer (Traugott et al., 2013).

Stable isotope analysis is one of the most valuable tool for studying food webs (Ehleringer et al., 1986; Fry, 2006; Boecklen et al., 2011; Traugott et al., 2013) and has been adopted for long to also study soil food webs (Ponsard and Arditi, 2000; Scheu and Falca, 2000;

Scheu, 2002). One of the disadvantages of this method is the rather large amount of material needed hampering the analysis of small species, this, however, has been overcome in part by recent advances in the analytical procedure (Langel and Dyckmans, 2014).

(26)

16

Natural variations in ¹⁵N/¹⁴N and ¹³C/¹²C ratios allow to evaluate the trophic structure of animal communities (DeNiro and Epstein, 1978; Minagawa and Wada, 1984; Wada et al., 1991). Stable isotope signatures of animal tissue provide information about the trophic position and trophic links of animals as well as on the basal resource used (Tiunov, 2007;

Traugott et al., 2013). The use of stable isotopes for analyzing the structure of food webs started in the 1970s (DeNiro and Epstein, 1978) and increased ever since (Ehleringer et al., 1986; Fry, 2006; Boecklen et al., 2011). ¹³C ratios are used to trace basal food resources, since ¹³C/¹²C ratios stay rather constant through food chains (Post, 2002). By contrast, isotopic fractioning leads to an enrichment in ¹⁵N from prey to consumer by 3.4 ‰ per trophic level (Post, 2002; Martínez Del Rio et al., 2009), thereby ¹⁵N values allow ascribing animals to different trophic levels and reflecting their feeding habits (DeNiro and Epstein, 1978; Kreipe et al., 2015). However, estimation of the trophic position requires an isotopic baseline since ¹⁵N/¹⁴N and ¹³C/¹²C ratios vary in primary producers in time and space (Jardine et al., 2006). To establish the appropriate baseline, it is necessary to measure the isotopic signatures of carbon sources within the study site, i.e. by measuring the stable isotope ratio of leaf litter and soil (Casey and Post, 2011). Comparing isotope signatures of animals and the basal resource is a powerful tool to understand trophic interactions and dynamics of organic matter in soil (Potapov et al., 2019), processes otherwise very difficult to study in-situ (Tiunov, 2007). Dead plant material either is used for decomposition or incorporated into soil organic matter, processes which are essentially driven by the activity and/or interactions between soil organisms (De Ruiter et al., 1993;

Nielsen et al., 2011; Filser et al., 2016; Potapov et al., 2019).

(27)

17

Stable isotopes cannot only be used to study the structure of food webs, but also to analyze the trophic ecology of specific taxonomic groups (Scheu and Falca, 2000; Halaj et al., 2005; Maraun et al., 2007, 2011; Tiunov, 2007). In soil, stable isotopes have been used to investigate trophic niches of species of earthworms, (Martin et al., 1992), ants (Blüthgen et al., 2003), but also smaller animals such as oribatid mites (Schneider et al., 2004; Erdmann et al., 2007; Maraun et al., 2011), springtails (Chahartaghi et al., 2005) and mesostigmatid mites (Klarner et al., 2013). Blüthgen et al. (2003) showed that stable isotopes analysis is a powerful tool for investigating trophic niche partitioning and plasticity in complex and diverse communities.

Management of oil palm plantations and the ‘Biodiversity Enrichment Experiment‘

Natural ecosystems are affected by human activities, including conversion into plantations and management of ecosystems (Foley et al., 2005, 2011). Until today, about 40 % of the terrestrial surface has been transformed into agricultural systems (Foley et al., 2011). In part, however, these systems are not managed in a sustainable way. Rather, a considerable proportion of them is degraded; further habitat loss occurs due to construction of infrastructure and desertification processes (Bridges and Oldeman, 1999;

Reynolds et al., 2007; Foley et al., 2011; Pavao-Zuckerman and Sookhdeo, 2017; Francini et al., 2018).

Intensification of land use and the associated biodiversity loss affects the structure of ecological communities and therefore the functioning of above- and belowground systems (Sodhi et al., 2004; Erdmann et al., 2007; Wilcove et al., 2013; Barnes et al., 2014;

(28)

18

Edwards et al., 2014; Klarner et al., 2017). Reduced decomposer diversity may compromise decomposition processes as well as carbon and nutrient cycling (Handa et al., 2014). The large scale transformation of rainforests into monoculture plantation systems, such as oil palm and rubber, is one of the main drivers for biodiversity loss, especially in South East Asia (Fitzherbert et al., 2008; Immerzeel et al., 2014; Teuscher et al., 2016).

Effects of oil palm plantation management on aboveground biodiversity and ecosystem functioning is receiving increased interest (Nurdiansyah et al., 2016; Syafiq et al., 2016;

Teuscher et al., 2016; Ashton-Butt et al., 2018), however, only few studies focused on consequences for the belowground systems (Bessou et al., 2017). As soil communities are linked to the diversity and abundance of plant communities (Eisenhauer et al., 2011;

Thakur and Eisenhauer, 2015), conversion of rainforest into monoculture plantations is likely to strongly affect belowground biodiversity. Therefore, to protect biodiversity of tropical regions it is important to integrate the belowground system and to consider the management of plantation systems (Koh et al., 2009; Foster et al., 2011; Luskin and Potts, 2011; Teuscher et al., 2015, 2016). Oil palm plantations may harbor a diverse understory (Foster et al., 2011), however, understory plants often compete with oil palms and therefore commonly are weeded (Tohiran et al., 2017). Removal of understory plants in oil palm plantations may be done by hand, but more commonly by the use of herbicides.

However, the extensive use of herbicides may pollute water and thus provides a threat to the already endangered ecosystems (Schiesari and Grillitsch, 2011; Comte et al., 2012). In fact, the use of pesticides in agricultural land-use systems has been linked to the decline in biodiversity (Geiger et al., 2010; Beketov et al., 2013). A number of studies showed that the reduction in herbicide use and the associated increase in the coverage of understory vegetation in oil palm plantations may beneficially affect aboveground invertebrates, but

(29)

19

also the decomposer system (Chung et al., 2000; Ashraf et al., 2018; Ashton-Butt et al., 2018; Spear et al., 2018; Darras et al., 2019).

(30)

20

Study site

Studies of the presented theis formed part of the interdisciplinary project “Ecological and socioeconomic functions of tropical lowland rainforest transformation sytems (Sumatra, Indonesia) (Drescher et al., 2016) (Fig. 1).

Figure 1. Location of study sites of EFForTS in Sumatra and the Jambi Province (Drescher et al., 2016), core plots are located in the two landscapes near Bukit Duabelas National Park and Harapan Rainforest.

(31)

21

Two (Chapters 2 and 3) of the three studies reported were conducted at the EFForTS core plots established 2012 in two landscapes, Bukit Duabelas (2° 0’ 57” S, 120° 45’ 12” E) and Harapan (1° 55’ 40” S, 103° 15’ 33” E). The dominant soil type at both landscapes is Acrisol.

At Bukit Duabelas soils with a clay texture predominate whereas Harapan soils are characterized by a sandy loam texture. In total, 32 core plots were established, four plots in each of four different land-use systems: lowland rainforest, jungle rubber, rubber and oil palm monoculture plantations. Rainforest plots represented “primary degraded forest”

(classified by Margono et al., 2014), with signs of selective logging as well as extraction of non-timber products. Jungle rubber, represented smallholder rubber agroforest systems comprising previously logged rainforest enriched with rubber trees (Hevea brasiliensis).

Both, rubber as well as oil palm plots were situated within smallholder monoculture plantations, varying between 7 to 16 years (rubber) and 8 to 15 years (oil palm) in 2012 (Drescher et al., 2016). Each plot was 50 x 50 m and contained 5 x 5 m subplots at random positions within the plot (Drescher et al., 2016). The third study (Chapter 4) formed part of the “Biodiversity Enrichment Experiment” established in 2013 in the framework of EFForTS in the oil palm plantation of PT Humusindo Makmur Sejata (01.95° S and 103.25°

E, 47±11 m a.s.l.) (Teuscher et al., 2016) (Fig. 2). The experiment was located in the Harapan landscape and the dominant soil type is loamy Acrisol (Allen et al., 2015). In 2016, the average age of planted oil palms was between 6 and 12 years. Management of the plantation contained fertilizer application, regular manual weeding of understory plants as well as removal of epiphytes. Herbicides were only applied if the manual weeding could not conducted due to lack of available workers (Teuscher et al., 2016).

(32)

22

Figure 2. Map of the study area of the EFForTS Project (Drescher et al., 2016, modified from Teuscher et al., 2016). The green star indicates the location of the enrichment experiment in the oil palm plantation of PT Humusindo Makmur Sejata.

Within the oil palm plantation, tree islands of varying species diversity and compositions were established (Fig. 3). 52 plots of different plot size (5 × 5, 10 × 10, 20 × 20 and 40 × 40 m) as well as different tree diversity levels (0, 1, 2, 3 and 6 species) were established, according to the random partitions design of Bell et al. (2009). Each tree species was selected only once at each species diversity level, therefore species composition within the tree islands was random. Additionally, four plots of the same size (10 x 10 m) and

(33)

23

management as usual were established as control plots (ctrl) (Teuscher et al., 2016), resulting in 56 plots total. For the enrichment of the tree islands six native trees were selected, i.e. three fruit trees (Parkia speciosa, and Archidendron pauciflorum, Fabaceae;

Durio zibethinus, Malvaceae), and three timber trees (Peronema canescens, Lamiaceae;

Shorea leprosula, Dipterocarpaceae), and one known to produces natural latex (Dyera polyphylla, Apocynaceae). Prior to tree planting, 40 % of the oil palms were removed from the experimental plots. Management of the established plots contained manual weeding in the first two years (preventing weeds to overgrow planted saplings; every three month) which was stopped after that to allow natural succession, i.e. interaction/competition with each other and oil palms. Application of fertilizers, herbicides and insecticides in the plots were stopped after tree planting. Samples for my thesis were taken in 2016 and therefore were without manual weeding.

(34)

24

Figure 3. Study design of the biodiversity experiment (Teuscher et al., 2016). (A) Tree island with

varying tree diversity levels (0, 1, 2, 3 and 6 species), identity and composition and plot size (5 × 5, 10 × 10, 20 × 20 and 40 × 40 m). Four control plots without treatment and with management as usual are represented by ctrl. In total, there are 56 plots. (B) Oil palms (OP) were cut in the plots to enhance the light conditions, planted trees are in a 2 x 2 m grid. (C) Planted trees interact/compete with each other and the oil palms.

Objectives and chapter outline

This thesis aims at improving the knowledge about the effects of land-use changes in South East Asia, Sumatra (Indonesia), i.e. more specifically the conversion of tropical rainforest to jungle rubber, rubber and oil palm monoculture plantations, on soil arthropod communities, especially oribatid mites, and their trophic ecology. Chapters 2 to 4 report results from field experiments with Chapter 2 investigating the shift in trophic

(35)

25

niches of individual species of oribatid mites with the conversion of tropical rainforest into plantations, as indicated by stable isotopes (¹⁵N, ¹³C). Chapter 3 investigates the shift in trophic niches of oribatid mite communities, represented by the species making up 80 % of total oribatid mite individuals in the respective land-use system, as indicated by stable isotopes (¹⁵N, ¹³C). Chapter 4 investigates the effect of the enrichment of oil palm plantations with native tree species in ‘tree islands’ as well as varying island size on soil invertebrate communities as part of the ‘Biodiversity Enrichment Experiment’.

The main hypotheses of this thesis are as follows:

(1) Oribatid mite species cope with environmental changes in transformed ecosystems by shifting their trophic niches, with land-use system change inducing a shift in trophic levels and/or the use of basal resources indicating trophic plasticity.

(2) The trophic niche of oribatid mite communities changes with land-use system being larger in more natural systems (rainforest, jungle rubber) than in plantation systems (rubber, oil palm).

(3) The enrichment of oil palm plantations with native tree species increases the density and complexity of soil arthropod communities with the effect increasing with plot size.

The content of the three chapters can be summarized as follows:

In Chapter 2 we investigated shifts in trophic niches of six soil-living oribatid mite species (Plonaphacarus kugohi, Protoribates paracapucinus, Scheloribates praeincisus, Bischeloribates mahunkai, Rostrozetes cf. shibai, and Rostrozetes sp. 1) with the conversion of lowland secondary rainforest into plantation systems of different land use

(36)

26

intensity (jungle rubber, rubber and oil palm monoculture plantation) in two regions of southwest Sumatra, Indonesia. We measured stable isotope ratios (¹³C/¹²C and ¹⁵N/¹⁴N) of single oribatid mite individuals and inspected shifts in stable isotope niches with changes in land-use systems. Significant shifts in stable isotope ratios in three of the six studied oribatid mite species (S. praeincisus, R. cf. shibai and Rostrozetes sp. 1) indicated that these species in fact shift their trophic niches with environmental changes. The trophic niche of the other three studied species (B. mahunkai, P. kugohi and P.

paracapucinus) did not differ significantly between the land-use systems, but generally followed a similar trend as in the other three species. Overall, the results suggest that colonization of very different ecosystems such as rainforest and intensively managed monoculture plantations by oribatid mite species likely is related to their ability to shift their trophic niches, i.e. to trophic plasticity. Notably, the shift was due to both changes in the use of basal resources as well as trophic levels.

Chapter 3 investigated shifts in the community-level trophic niche of oribatid mites with the conversion of rainforest into rubber and oil palm plantations. We investigated 80 % of oribatid mite communities occurring in lowland secondary rainforest and plantation systems of different land use intensity (jungle rubber, rubber and oil palm monoculture plantation) in two regions of southwest Sumatra, Indonesia. We measured stable isotope ratios (¹³C/¹²C and ¹⁵N/¹⁴N) of pooled individuals of oribatid mite species and inspected shifts in community-level trophic niche with changes in land-use systems. Our results confirmed that the community-level trophic niche of oribatid mites in fact is wider in rainforest than in plantation systems. Between natural and plantation systems a clear separation of the community-level trophic niche occur, indicating that with natural and

(37)

27

plantation systems the community-level trophic niche of oribatid mites is totally different.

As indicated by minimum and maximum of litter-calibrated isotopic signatures of oribatid mite community-level trophic niche, only oribatid mite isotopic signatures from oil palm or rubber were significantly different compared to rainforest and jungle rubber. This implies that within oil palm and rubber plantations, there are single species within oribatid mite communities which occupy trophic niches which are not present in rainforest and/or jungle rubber.

The study reported in Chapter 4 was part of the ‘Biodiversity Enrichment Experiment’ (see above). This experiment aimed at enhancing biodiversity and ecosystem functioning in oil palm plantations via ‘tree islands’ with varying diversity level (0, 1, 2, 3 and 6 different tree species) and plot size (5 x 5, 10 x 10, 20 x 20 and 40 x 40 m). We investigated the effect of

‘tree islands’ on macro- and mesofauna soil invertebrate taxa three years after the experiment was established. Our results demonstrated that neither the diversity level of the planted tree species nor plot size affected the abundance of soil invertebrate taxa but soil invertebrate richness varied with tree diversity. Notably, richness of soil invertebrates peaked at diversity level 2. As soil communities respond with a delay in time to soil forming process, we expect that the observed changes will increase in time.

References

Abdullah, S.A., Nobukazu, N., 2007. Forest fragmentation and its correlation to human land use change in the state of Selangor, peninsular Malaysia. Forest Ecology and Management 241.1-3, 39–48.

Abrams, P.A., 1987. On classifying interactions between populations. Oecologia 73, 272–281.

Ahrends, A., Hollingsworth, P.M., Ziegler, A.D., Fox, J.M., Chen, H., Su, Y., Jianchu, X., 2015. Current trends of rubber plantation expansion may threaten biodiversity and livelihoods. Global Environmental Change 34, 48–58.

(38)

28

Allen, K., Corre, M.D., Tjoa, A., Veldkamp, E., 2015. Soil nitrogen-cycling responses to conversion of lowland forests to oil palm and rubber plantations in Sumatra, Indonesia. PLOS ONE 10, 1–21.

Almeida, D., Almodóvar, A., Nicola, G.G., Elvira, B., Grossman, G.D., 2012. Trophic plasticity of invasive juvenile largemouth bass Micropterus salmoides in Iberian streams. Fisheries Research 113, 153–158.

Ashraf, M., Zulkifli, R., Sanusi, R., Tohiran, K.A., Terhem, R., Moslim, R., Norhisham, A.R., Ashton-Butt, A., Azhar, B., 2018. Alley-cropping system can boost arthropod biodiversity and ecosystem functions in oil palm. Agriculture, Ecosystems & Environment 260, 19–26.

Ashton, L.A., Griffiths, H.M., Parr, C.L., Evans, T.A., Didham, R.K., Hasan, F., Teh, Y.A., Tin, H.S., Vairappan, C.S., Eggleto, P., 2019. Termites mitigate the effects of drought in tropical rainforest. Science 363, 174–177.

Ashton-Butt, A., Aryawan, A.A., Hood, A.S., Naim, M., Purnomo, D., Suhardi, S., Wahyuningsih, R., Willcock, S., Poppy, G.M., Caliman, J., Turner, E.C., Foster, W., Peh, K.S., Snaddon, J.L., 2018.

Understory vegetation in oil palm plantations benefits soil biodiversity and decomposition rates, Frontier in Forest and Global Change.

Bardgett, R., 2005. The biology of soil: A community and ecosystem approach. Oxford University Press.

Barnes, A.D., Jochum, M., Mumme, S., Haneda, N.F., Farajallah, A., Widarto, T.H., Brose, U., 2014.

Consequences of tropical land use for multitrophic biodiversity and ecosystem functioning. Nature Communications 5, 1–7.

Bedano, J.C., Domínguez, A., Arolfo, R., 2011. Assessment of soil biological degradation using mesofauna.

Soil and Tillage Research 117, 55–60.

Behan-Pelletier, V.M., 1999. Oribatid mite biodiversity in agroecosystems: Role for bioindication.

Agriculture, Ecosystems and Environment 74, 411–423.

Beketov, M. A., Kefford, B.J., Schafer, R.B., Liess, M., 2013. Pesticides reduce regional biodiversity of stream invertebrates. Proceedings of the National Academy of Sciences 110, 11039–11043.

Bell, T., Lilley, A.K., Hector, A., Schmid, B., King, L., Newman, J.A., 2009. A linear model method for biodiversity-ecosystem functioning experiments. The American Naturalist 174, 836–849.

Berg, B., McClaugherty, C., 2008. Plant Litter: Decomposition, humus formation carbon sequestration.

Springer-Verlag Heidelberg.

Berg, M.P., Bengtsson, J., 2007. Temporal and spatial variability in soil food web structure. Oikos 116, 1789–

1804.

Bessou, C., Verwilghen, A., Beaudoin-Ollivier, L., Marichal, R., Ollivier, J., Baron, V., Bonneau, X., Carron, M.-P., Snoeck, D., Naim, M., Ketuk Aryawan, A.A., Raoul, F., Giraudoux, P., Surya, E., Sihombing, E., Caliman, J.-P., 2017. Agroecological practices in oil palm plantations: Examples from the field. Ocl 24.

Beukema, H., Danielsen, F., Vincent, G., Hardiwinoto, S., Van Andel, J., 2007. Plant and bird diversity in rubber agroforests in the lowlands of Sumatra, Indonesia. Agroforestry Systems 70, 217–242.

(39)

29

Birch, L.C., 1957. The meanings of competition. The American Naturalist 91, 5–18.

Blüthgen, N., Gebauer, G., Fiedler, K., 2003. Disentangling a rainforest food web using stable isotopes:

Dietary diversity in a species-rich ant community. Oecologia 137, 426–435.

Boecklen, W.J., Yarnes, C.T., Cook, B.A., James, A.C., 2011. On the use of stable isotopes in trophic ecology.

Annual Review of Ecology, Evolution, and Systematics 42, 411–440.

Bommarco, R., Biesmeijer, J.C., Meyer, B., Potts, S.G., Pöyry, J., Roberts, S.P.M., Steffan-Dewenter, I., Ockinger, E., 2010. Dispersal capacity and diet breadth modify the response of wild bees to habitat loss. Proceedings of the Royal Society B: Biological Sciences 277, 2075–2082.

Bonachela, J.A., Pringle, R.M., Sheffer, E., Coverdale, T.C., Guyton, J.A., Caylor, K.K., Levin, S.A., Tarnita, C.E., 2015. Termite mounds can increase the robustness of dryland ecosystems to climatic change.

Science 347, 651–655.

Bowen, S.H., Allanson, B.R., 1982. Behavioral and trophic plasticity of juvenile Tilapia mossambica in utilization of the unstable littoral habitat. Environmental Biology of Fishes 7, 357–362.

Bridges, E.M., Oldeman, L.R., 1999. Global assessment of human-induced soil degradation. Arid Soil Research and Rehabilitation 13, 319–325.

Brückner, A., Schuster, R., Smit, T., Heethoff, M., 2018. Imprinted or innate food preferences in the model mite Archegozetes. Soil Organisms 90, 23–26.

Brussaard, L., Pulleman, M.M., Ouédraogo, É., Mando, A., Six, J., 2007. Soil fauna and soil function in the fabric of the food web. Pedobiologia 50, 447–462.

Carter, C., Finley, W., Fry, J., Jackson, D., Willis, L., 2007. Palm oil markets and future supply. European Journal of Lipid Science and Technology 109, 307–314.

Casey, M.M., Post, D.M., 2011. The problem of isotopic baseline: Reconstructing the diet and trophic position of fossil animals. Earth-Science Reviews 106, 131–148.

Chahartaghi, M., Langel, R., Scheu, S., Ruess, L., 2005. Feeding guilds in Collembola based on nitrogen stable isotope ratios. Soil Biology and Biochemistry 37, 1718–1725.

Chase, J.M., Leibold, M.A., 2003. Ecological niches: Linking classical and contemporary approaches.

University of Chicago Press.

Chung, A.Y.C., Eggleton, P., Speight, M.R., Hammond, P.M., Chey, V.K., 2000. The diversity of beetle assemblages in different habitat types in Sabah, Malaysia. Bulletin of Entomological Research 90, 475–

496.

Clay, J., 2013. World agriculture and the environment: A commodity-by-commodity guide to impacts and practices. Island Press.

(40)

30

Comte, I., Colin, F., Whalen, J.K., Grünberger, O., Caliman, J.P., 2012. Agricultural practices in oil palm plantations and their impact on hydrological changes, nutrient fluxes and water quality in Indonesia.

A review., 1st ed, Advances in Agronomy. Elsevier Inc.

Corley, R., Hereward, V., Tinker, P.B., 2008. The oil palm. John Wiley & Sons.

Coûteaux, M.M., Bottner, P., Berg, B., 1995. Litter decomposition, climate and liter quality. Trends in Ecology & Evolution 12.2, 63–66.

Curran, L.M., Trigg, S.N., Mcdonald, A.K., Astiani, D., Hardiono, Y.M., Siregar, P., Caniago, I., Kasischke, E., 2004. Lowland forest loss in protected areas. Science 303, 1000–1003.

Dammhahn, M., Randriamoria, T.M., Goodman, S.M., 2017. Broad and flexible stable isotope niches in invasive non-native Rattus spp. in anthropogenic and natural habitats of central eastern Madagascar.

BMC Ecology 17, 1–13.

Danielsen, F., Beukema, H., Burgess, N.D., Parish, F., Brühl, C.A., Donald, P.F., Murdiyarso, D., Phalan, B., Reijnders, L., Struebig, M., Fitzherbert, E.B., 2009. Biofuel plantations on forested lands: Double jeopardy for biodiversity and climate. Conservation Biology 23, 348–358.

Darras, K., Corre, M.D., Formaglio, G., Tjoa, A., Potapov, A., Sibhatu, K.T., Grass, I., Tscharntke, T., Rubiano, A.A., Buchori, D., Drescher, J., Fardiansah, R., Hölscher, D., Irawan, B., Kneib, T., Krashevska, V., Krause, A., Kreft, H., Li, K., Polle, A., Ryadin, A.R., Rembold, K., Scheu, S., 2019. Reducing fertilizer and avoiding herbicides in oil palm plantations - Ecological and economic valuation. Frontiers in Forests and Global Change 2, 65.

David, J.F., 2014. The role of litter-feeding macroarthropods in decomposition processes: A reappraisal of common views. Soil Biology and Biochemistry 76, 109–118.

De Ruiter, P.C., Van Veen, J.A., Moore, J.C., Brussaard, L., Hunt, H.W., 1993. Calculation of nitrogen mineralization in soil food webs. Plant and Soil 157, 263–273.

De Vries, F.T., Thebault, E., Liiri, M., Birkhofer, K., Tsiafouli, M. a., Bjornlund, L., Bracht Jorgensen, H., Brady, M. V., Christensen, S., de Ruiter, P.C., d’Hertefeldt, T., Frouz, J., Hedlund, K., Hemerik, L., Hol, W.H.G., Hotes, S., Mortimer, S.R., Setala, H., Sgardelis, S.P., Uteseny, K., van der Putten, W.H., Wolters, V., Bardgett, R.D., 2013. Soil food web properties explain ecosystem services across European land use systems. Proceedings of the National Academy of Sciences 110, 14296–14301.

Deagle, B.E., Kirkwood, R., Jarman, S.N., 2009. Analysis of Australian fur seal diet by pyrosequencing prey DNA in faeces. Molecular Ecology 18, 2022–2038.

DeFries, R., Foley, J., Asner, G., 2004. Balancing human needs and ecosystem function. Front Ecol Environ 2, 249–257.

DeNiro, M.J., Epstein, S., 1978. Influence of diet on the distribution of nitrogen isotopes in animals.

Geochimica et Cosmochimica Acta 45, 341–351.

Dennis, R. a., Mayer, J., Applegate, G., Chokkalingam, U., Colfer, C.J.P., Kurniawan, I., Lachowski, H., Maus, P., Permana, R.P., Ruchiat, Y., Stolle, F., Suyanto, Tomich, T.P., 2005. Fire, people and pixels: Linking

(41)

31

social science and remote sensing to understand underlying causes and impacts of fires in Indonesia.

Human Ecology 33, 465–504.

Di Lascio, A., Rossi, L., Pasquale, C., Calizza, E., Rossi, D., Costantini, M.L., Achtziger, R., 2013. Stable isotope variation in macroinvertebrates indicates anthropogenic disturbance along an urban stretch of the river Tiber (Rome, Italy). Ecological Indicators 28, 107–114.

Dirzo, R., Raven, P.H., 2003. Global state of biodiversity and loss. Annual Review of Environment and Resources 28, 137–167.

Donald, P.F., 2004. Biodiversity impacts of some agricultural. Issues Int. Conserv. 18, 17–37.

Drescher, J., Rembold, K., Allen, K., Beckscha, P., Buchori, D., Clough, Y., Faust, H., Fauzi, A.M., Gunawan, D., Hertel, D., Irawan, B., Jaya, I.N.S., Klarner, B., Kleinn, C., Knohl, A., Kotowska, M.M., Krashevska, V., Krishna, V., Leuschner, C., Lorenz, W., Meijide, A., Melati, D., Steinebach, S., Tjoa, A., Tscharntke, T., Wick, B., Wiegand, K., Kreft, H., Scheu, S., 2016. Ecological and socio-economic functions across tropical land use systems after rainforest conversion. Philosophical Transactions of the Royal Society of London. Series B, Biological Sciences 231, 1–7.

Drymon, J.M., Powers, S.P., Carmichael, R.H., 2012. Trophic plasticity in the Atlantic sharpnose shark (Rhizoprionodon terraenovae) from the north central Gulf of Mexico. Environmental Biology of Fishes 95, 21–35.

Ducarme, X., André, H.M., Wauthy, G., Lebrun, P., 2004. Are there real endogeic species in temperate forest mites? Pedobiologia 48, 139–147.

Edwards, F.A., Edwards, D.P., Larsen, T.H., Hsu, W.W., Benedick, S., Chung, a., Vun Khen, C., Wilcove, D.S., Hamer, K.C., 2014. Does logging and forest conversion to oil palm agriculture alter functional diversity in a biodiversity hotspot? Animal Conservation 17, 163–173.

Ehleringer, J.R., Rundel, P.W., Nagy, K. a., 1986. Stable isotopes in physiological ecology and food web research. Trends in Ecology and Evolution 1, 42–45.

Eisenhauer, N., Milcu, A., Sabais, A.C.W., Bessler, H., Brenner, J., Engels, C., Klarner, B., Maraun, M., Partsch, S., Roscher, C., Schonert, F., Temperton, V.M., Thomisch, K., Weigelt, A., Weisser, W.W., Scheu, S., 2011. Plant diversity surpasses plant functional groups and plant productivity as driver of soil biota in the long term. PLOS ONE 6, 15–18.

Elton, C.S., 1927. Animal ecology: London. Sidgwick and Jackson, Ltd.

Erdmann, G., Otte, V., Langel, R., Scheu, S., Maraun, M., 2007. The trophic structure of bark-living oribatid mite communities analysed with stable isotopes (¹⁵N,¹³C) indicates strong niche differentiation.

Experimental and Applied Acarology 41, 1–10.

Ettema, C.H., Wardle, D.A., 2002. Spatial soil ecology. Trends in Ecology & Evolution 17.4, 177–183.

European Comission, 2006. An EU Strategy for biofuels, impact assessment. 18.